The Journal of Wildlife Management 79(6):898–906; 2015; DOI: 10.1002/jwmg.915
Research Article
Indirect Effects of Red Imported Fire Ants on
Attwater’s Prairie-Chicken Brood Survival
MICHAEL E. MORROW,1 Attwater Prairie Chicken National Wildlife Refuge, PO Box 519, Eagle Lake, TX 77434, USA
REBECCA E. CHESTER, Attwater Prairie Chicken National Wildlife Refuge, PO Box 519, Eagle Lake, TX 77434, USA
SARAH E. LEHNEN, US Fish and Wildlife Service, 500 Gold Avenue SW, Room 4005, Albuquerque, NM 87102, USA
BASTIAAN M. DREES, Texas A&M AgriLife Extension Service Department of Entomology, Texas A&M University, Minnie Belle Heep,
Room 412, College Station, TX 77843, USA
JOHN E. TOEPFER, Society of Tympanuchus Cupido Pinnatus, Ltd., Ada, MN 56510, USA
ABSTRACT The invasive red imported fire ant (Solenopsis invicta) has negatively affected a host of
taxonomic groups throughout its acquired North American range. Many studies have hypothesized indirect
trophic impacts, but few documented those impacts. We evaluated invertebrate abundance as a factor limiting
juvenile survival of the endangered Attwater’s prairie-chicken (Tympanuchus cupido attwateri), and whether
fire ants reduce invertebrate numbers and biomass. From 2009–2013, we monitored survival of Attwater’s
prairie-chicken broods (n ¼ 63) with radio telemetry during the first 2 weeks post-hatch and collected daily
invertebrate samples at brood sites. Broods located in areas with the highest median invertebrate count (338
invertebrates/25 sweeps) had a survival probability of 0.83 at 2 weeks post-hatch compared to 0.07 for broods
located in areas with the lowest median invertebrate count (18 invertebrates/25 sweeps). During 2011–2012,
we evaluated the reduction of fire ants on invertebrate numbers and biomass by aerially treating areas with
Extinguish PlusTM in an impact-reference study design. Treated fields had 27% more individual invertebrates
and 26% higher invertebrate biomass than reference fields. Our results clearly document that invertebrate
abundance affects Attwater’s prairie-chicken brood survival and that fire ants may indirectly contribute to low
brood survival by suppressing invertebrate abundance. We posit that within the fire ant’s acquired North
American range, fire ants are likely contributing to declines of other insectivorous species. Published 2015.
This article is a U.S. Government work and is in the public domain in the USA.
KEY WORDS Attwater’s, brood survival, fire ants, indirect effects, invertebrates, prairie-chicken, Tympanuchus cupido
attwateri.
Red imported fire ants (Solenopsis invicta) were accidentally
introduced to the United States at the port of Mobile,
Alabama, circa 1930 (Allen et al. 1995), and arrived in
Attwater’s prairie-chicken (Tympanuchus cupido attwateri;
hereafter Attwater’s) range circa 1970 (Allen et al. 1995,
eXtension 2008). Today these fire ants occupy almost the
entire southeastern United States (Porter and Savignano
1990). Fire ants affect a host of fauna, including insects (e.g.,
Porter and Savignano 1990, Drees 1994, Mueller et al. 1999,
Wojcik et al. 2001, Allen et al. 2004). Although there is
considerable evidence documenting direct adverse effects at
the individual level, few studies address population-level
impacts, and fewer still provide experimental evidence
regarding functional mechanisms of indirect impacts.
Received: 15 July 2014; Accepted: 7 May 2015
Published: 17 June 2015
1
E-mail: [email protected]
This is an open access article under the terms of the Creative Commons
Attribution-NonCommercial-NoDerivs License, which permits use and
distribution in any medium, provided the original work is properly cited,
the use is non-commercial and no modifications or adaptations are made.
898
The Attwater’s prairie-chicken is endemic to prairies of the
Texas and Louisiana gulf coast (Bendire 1894), and was
listed as an endangered species in 1967 (U.S. Fish and
Wildlife Service [USFWS] 2010). Historically approaching
1 million individuals on 2.4 million ha of habitat (Lehmann
1968), fewer than 110 have existed in the wild since 1995
despite intensive management (USFWS 2010). Recovery
actions undertaken during the last 20 years have included
habitat restoration, captive breeding and release, and a
number of research projects focused on identifying factors
limiting recovery progress (USFWS 2010). The Attwater’s
Prairie-Chicken Recovery Plan identified poor survival of
chicks in the wild as “. . . the single-most important factor
limiting significant progress toward recovery” (USFWS
2010:40). Observations at the Attwater Prairie Chicken
National Wildlife Refuge (APCNWR) in 2003 revealed the
total loss of 8 broods <12 days post-hatch. Several dead or
dying chicks were found with brood hens at night roosts
(USFWS 2010). Necropsy of those chicks attributed cause of
death to inanition and dehydration.
Invertebrates are a critical food source for young prairiechickens, including Attwater’s (e.g., Lehmann 1941, Jones
The Journal of Wildlife Management
79(6)
1963, Rumble et al. 1988, Savory 1989, Hagen et al. 2005).
Savory (1989) noted that invertebrate food comprised
>90% of the diet of T. cupido <5–8 weeks old. From 2004–
2012, we confined broods (i.e., hen and chicks) at nest sites
for 2 weeks post-hatch and provided locally collected
insects ad libitum. Most (83% of 547) chicks treated in this
manner survived the critical 2-week period. These
observations led us to hypothesize that reduced invertebrate
abundance resulting from fire ant invasion was limiting
Attwater’s chick survival. In this study, we investigated
the relationship between invertebrate abundance and
Attwater’s brood survival and the effects of fire ants on
invertebrate abundance at multiple large-scale treatment
sites during the Attwater’s early brooding season (May–
mid-June).
STUDY AREA
We conducted our study at 5 locations in Texas within the
Attwater’s prairie-chicken’s current or historical range. The
970 ha Texas City Prairie Preserve (TCPP; Galveston
County) was owned and managed by The Nature
Conservancy of Texas. The 4,265-ha APCNWR (Colorado
County) was part of the USFWS National Wildlife Refuge
System, and was established specifically to maintain habitat
for Attwater’s prairie-chickens. The Midwell (Goliad
County), West Clarkson (Goliad County), and Tivoli
(Refugio County) study sites were located on private cattle
ranches situated within approximately 20,445 ha of relatively
contiguous grasslands (USFWS 2010). All 5 sites were
within the gulf prairies and marshes vegetational area of
Texas (Hatch et al. 1990) and consisted of managed open
grasslands dominated by mid–tall native grass species. Four
of the sites had wild Attwater’s populations, which had been
supplemented with captive-reared birds (Morrow et al. 2004,
USFWS 2010).
Climate of the region was subtropical, and dominated by
warm, moist air masses derived from the Gulf of Mexico
(Smeins et al. 1991). Annual precipitation ranged from
approximately 864 mm at the western-most site (Midwell) to
1,118 mm for the eastern-most site (TCPP; Hatch et al.
1990). During 2011, extreme drought conditions prevailed
throughout the region, and rainfall totals were only 41–57%
of long-term averages for our study areas (National Oceanic
Atmospheric Association 2013).
All sites were grazed by cattle during our study. Grazing at
APCNWR and Midwell was at light to moderate levels,
West Clarkson and Tivoli received moderate to heavy
grazing, and TCPP received heavy grazing (M. E. Morrow,
Attwater Prairie Chicken National Wildlife Refuge,
unpublished data; C. Carter, Ranch Consultant, and A.
Tjelmeland, The Nature Conservancy, personal communication). Grazing pressure was exacerbated during drought
conditions. All sites used prescribed burning as a management tool except Tivoli, which had no history of prescribed
burning for more than 25 years. The Midwell and West
Clarkson sites were aerially treated in a strip pattern (approx.
75–90% coverage) by ranch managers with a broadleaf
herbicide (2,4-D) in late April–early May of 2012 to suppress
Morrow et al.
Fire Ants and Attwater’s Prairie-Chicken
forbs and minimize competition with forage grasses (C.
Carter, personal communication).
METHODS
We treated a 308-ha block at APCNWR by helicopter with
1.7 kg/ha (1.5 lbs/ac) of Extinguish PlusTM (0.365%
hydramethylnon, 0.25% s-methoprene; Central Life Sciences, Schaumburg, IL) fire ant bait to suppress fire ants as a
pilot project during April and November 2009. Bait products
such as Extingush Plus are specific to ants, and have few
environmental risks compared to contact insecticides (Barr
et al. 2005, Knutson and Campos 2008). Although other ant
species are susceptible to this product, fire ants dominate the
insecticide-laden bait product to their detriment while
allowing native ant population increases (Calixto et al. 2007).
During November 2010 and again in late-September 2011,
we treated 527, 294, 204, 259, and 179 ha similarly at
APCNWR, Midwell, Texas City, Tivoli, and West
Clarkson, respectively. We treated only APCNWR in
September 2012. We applied Extinguish Plus during
weather conditions appropriate for application (no moisture
on vegetation, temperature >218C). We made on-theground observations during application to verify that the bait
was picked up by fire ants.
Brood Survival
From 2009–2013, we evaluated the relationship between
brood survival and invertebrate abundance at the APCNWR
and Midwell study sites by monitoring wild Attwater’s
broods. Attwater’s hens were equipped with ponchomounted radio transmitters with tuned-loop antennas
(Telemetry Solutions, Walnut Creek, CA; Advanced
Telemetry Systems, Isanti, MN) incorporated into the
poncho (Amstrup 1980, Toepfer 2003). A limited number of
transmitters (American Wildlife Enterprises, Monticello,
FL) were equipped with whip antennas. We trimmed
antennas on these transmitters to extend only approximately
6 cm beyond the poncho to avoid potential interference with
flight (Marks and Marks 1987). We placed transmitters
(3% of body mass) on hens several months in advance of
nesting.
We tracked broods (hen with chicks) using telemetry
approximately 10 times (i.e., daily during the 5-day work
week) during the first 2 weeks post-hatch, except when
adverse weather hindered obtaining these data or hen
movements (successive >1.6 km) indicated probable brood
loss. We determined brood hen locations either by
triangulation using a vehicle-mounted 6-element yagi
antenna system and computer software (DogTrack, Blacksburg, VA; Locate III, Tatamagouche, NS, Canada), or by
circling brood hens at a distance of 10–20 m with a handheld telemetry system and recording location data with a
global positioning system (GPS) unit. We classified brood
locations with respect to fire ant treatment (i.e., located in an
area treated within the past year or not), and determined the
proportion (%) of locations occurring within treated areas.
Radiomarking of hens and brood monitoring activities were
authorized by Federal Fish and Wildlife Permit TE051839
899
and Texas Parks and Wildlife Department Scientific
Research Permit SPR-0491-384.
We collected invertebrate samples at brood locations 1–4
days later to avoid disturbance to the broods. We made 25
vigorous sweeps of vegetation along an approximately 20-m
transect in a random direction from the pre-determined
brood location with a 38-cm diameter, canvas sweep-net.
Each sweep consisted of a single approximately 2-m arc of
the net. We did not collect samples when vegetation was wet.
We labeled invertebrate samples and froze them until we
could determine counts of individuals for each sample. We
then dried samples for 24 hours at 60 8C and weighed them
to the nearest 0.0001 g on a digital analytical scale. We
determined median invertebrate numbers and dry mass for
each brood to control for pseudoreplication within brood
unit (Crawley 2007). We censored broods if the hen died
prior to 14 days post-hatch.
We approached radioed hens with hand-held telemetry
equipment 14 days post-hatch and observed them at dawn,
before leaving night roosts, to determine if any chicks
remained with the hen. We encouraged hens to move only if
necessary to be absolutely certain of chick presence. We
considered a brood successful if we observed 1 chick with
the hen. To reduce disturbance to the hen and brood, we did
not attempt to obtain a total count of chicks.
stored in alcohol. We sorted invertebrates into 2 groups to
differentiate between those that were more prone to flight
and therefore could have moved into the plot, from those
that were more typically ground-dwelling and less likely to
have immigrated from outside treatment areas. Invertebrates
more prone to flight included adult (indicated by wings
protruding behind the posterior tip of the abdomen)
grasshoppers or katydids (Orthoptera), Dipterans, adult
Lepidopterans, and bees or wasps (Hymenoptera). Grounddwelling invertebrates included juvenile grasshoppers or
katydids; leafhoppers (Cicadellidae); beetles, weevils, and
grubs (Coleoptera); crickets (Gryllidae); larval Lepidopterans, mantids or walking sticks (Mantodea, Phasmatodea);
ants (Formicidae); and snails (Gastropoda). These 2 groups
(hereafter mobile-flier and ground-dwelling, respectively)
also likely represented functional groups for Attwater’s
chicks because mobile-fliers would be more difficult for
chicks to catch than the ground-dwelling group. More than
85% of the invertebrate species found in Attwater’s diets by
Lehmann (1941) would likely have been classified as grounddwelling based on the season of collection. We removed 9
samples with missing or uncertain observations, resulting in a
total of 831 invertebrate samples for analysis (n ¼ 168 each
for Midwell, West Clarkson, and Tivoli; 166 for APCNWR;
and 161 for TCPP).
Impact of Fire Ant Suppression on Invertebrate
Abundance
During 2011–2012, we assessed invertebrate abundance and
biomass following fire ant suppression using an impactreference design replicated at 5 locations (APCNWR,
Midwell, TCPP, Tivoli, West Clarkson) with 3 temporal
replicates during each year. We randomly assigned treated
and untreated reference areas at each of the 5 sites. Each
treated area received 2 applications of Extinguish Plus
(autumn 2010 and 2011) as described previously. The pilot
project area at APCNWR was not included in this portion of
the study.
We sampled invertebrates and fire ant activity on each
location replicate at 14 0.2-ha randomly located plots within
treated and reference areas (28 sampling plots/site).
Sampling plots were >61 m from treatment edges, and
were stratified by ecological site and time since burn. We
sampled areas each year on 3 occasions separated by about
2 weeks starting in late April–early May following the
autumn applications of Extinguish Plus. We assessed fire ant
activity within each sampling plot by placing 10 fatty lures
(approx. 0.5-inch hot dog slices) on the ground using
surveyor’s flags, 1 in the center and 9 spaced roughly
equidistantly around the perimeter of the 0.2-ha plot. After
45–60 minutes, we estimated and recorded the number of fire
ants on the top and sides of each hot dog slice (Drees 2002).
We estimated numbers of ants present in increments of 10 up
to a maximum of 100. We obtained invertebrate samples by
sweep-netting in a random direction using the procedure
described previously for prairie-chicken brood sites and froze
samples for 24 hours. Invertebrates were separated into 2
groups, counted, dried, weighed to the nearest 0.0001 g, and
Data Analyses
Brood survival.—We determined the relationship of
invertebrate abundance (counts and biomass) and fire ant
treatment with brood survival (Sb) using generalized linear
mixed effects models. The response variable was presence or
absence of at least 1 chick at 2 weeks post-hatch. Explanatory
variables included the median count and biomass of
invertebrate samples collected during the first 2 weeks after
hatch for each brood, year, site, and the percentage of
observations for each brood that were in treated fields. We
used hen ID to specify a random effect (intercept). By
including hen as a random variable, we avoided pseudoreplication resulting from the inclusion of hens with more
than one brood in successive years. We built models using the
glmer function in the package lme4 (Bates et al. 2014) in the
R statistical system (R Version 3.2.0, www.r-project.org,
accessed 27 May 2015) and specified a binomial distribution
with a logit link. We identified top models using change in
Akaike’s Information Criterion values (DAIC; Akaike 1973)
and associated Akaike weights (w; Burnham and Anderson
2002). We report 90% rather than 95% confidence intervals
because 1) given that the Attwater’s prairie-chicken is a
critically endangered species, we considered the risk of
mistakenly concluding no treatment effect when there was
one (type II error) a worse error than the risk of concluding
there was a treatment effect when there was not (type I error);
and 2) the small size of the remaining population of
Attwater’s prairie-chickens meant that the brood sample size
was also small.
Effect of fire ant suppression on invertebrate abundance.—We
determined the effect of fire ant treatment on invertebrate
count and biomass using linear and generalized linear mixed
900
The Journal of Wildlife Management
79(6)
effects models. Random effects in these models accounted for
lack of independence among samples collected from the
same plots. Although differences among sites were not of
particular interest to us and therefore could have have been
included as random effects, we included site as a fixed effect
because sample size (n ¼ 5) was too small to include this
variable as a random effect (Bolker et al. 2009). The basic
procedure for model fitting was the same for both linear
and generalized linear mixed effects models. We first fit
competing models of random effects using the same global
model of fixed effects for all models (Zuur et al. 2009). We
compared 4 competing models of random effects: a random
intercept of the plot only model and 3 models with a random
intercept of plot and a random slope of plot across time (year,
day of sample, and sample period). The fixed effects in the
global model were year, site, treatment, and an interaction
between site and year. We used the random effects
component in the model with the lowest AIC in subsequent
models comparing fixed effects. Our candidate set of fixed
effects included 9 models using combinations of treatment,
site, year, the interaction between site and year, and the
interaction between treatment and year. We included the
interaction between treatment and year to account for
possible greater treatment effects in the second year, after
2 years of fire ant suppression. We identified top models
using DAIC and w values. We ran all analyses in the R
statistical system.
Response variables for count analyses included counts of
individuals for the mobile-flier and ground-dwelling
invertebrate groups, and for total invertebrates. Using
generalized models for count data is preferred because
they can model dispersion and eliminate the possibility of
predicting negative counts (O’Hara and Kotze 2010). For all
count data sets, variances of the counts were much larger
(range 2–72 for all site and year combinations) than the
mean; therefore, we used a negative binomial distribution to
account for this overdispersion. We visually inspected for
outliers and removed 4 (0.5%) deviant observations. We
compared data sets with outliers removed to unaltered data
sets to verify that removing outliers did not alter the
magnitude or direction of parameter estimates. We built
count data analysis models using the glmmadmb function in
the package glmmADMB (Fournier et al. 2012, Skaug et al.
2012).
For biomass data, we used linear mixed effects models for
ground-dwelling and total invertebrates. We Box-Cox
transformed data to approximate normality (Box and Cox
1964). We selected the type of transformation for each data
set using the powerTransform function in the car package
(Fox and Weisberg 2011). We visually inspected transformed data using quantile–quantile plots to verify normality
and did not detect any outliers. We built models using the
lmer function in the package lme4 (Bates et al. 2013).
Biomass data of mobile-flier invertebrates had an excess
amount of zeros and we could not find a transformation that
would approximate normality. Therefore, we used a hurdle
model for this data set (Cragg 1971). The log-normal hurdle
model consists of 2 parts. The first part is a binomial model
Morrow et al.
Fire Ants and Attwater’s Prairie-Chicken
that divides the data into its zero and non-zero components;
the second part models the positive range of the data set. This
model allows for the possibility that the mechanisms that
determine the zero and non-zero components can differ
(Ridout et al. 1998). In other words, the hurdle model tests
whether fire ant treatment affects the presence of mobile-flier
invertebrates within a field, and then separately, given that
mobile-fliers are present, tests the effect that fire ant
treatment has on mobile-flier biomass. We built hurdle
models using the glmmadmb function in the package
glmmADMB.
RESULTS
Brood Survival
We collected brood survival data from 36 hens with 44
broods at APCNWR and 17 hens with 19 broods at
Midwell. Of these 63 broods, 28 (44%) successfully
maintained chicks until 14 days post-hatch. Twenty-one
(33%) broods were in treatment areas at some point during
observation, although most broods remained entirely within
treatment (27%) or reference fields (67%). Of 427 brood
locations, 29% were in fields treated to suppress fire ants.
We made a median 7 observations on individual broods
(range 1–12).
The median number of total invertebrates at brood sites
was the best predictor of 2-week brood survival (Sb), with
6 times more support for this model than for the next best
model (Table 1; Fig. 1). The effect of median number of
invertebrates (b ¼ 1.45, 90% CI 0.43–2.41) on brood survival
was positive; broods located in areas with the highest median
invertebrate counts (338 invertebrates/sample) were more
than 11 times more likely to have chicks at 2 weeks posthatch (Sb ¼ 0.83, 90% CI 0.56–0.95) than those areas with
the lowest count (18 invertebrates/sample; Sb ¼ 0.07, 90% CI
0.01–0.31). With considerably less support, invertebrate
biomass (b ¼ 1.09, 90% CI 0.13–1.99) was the next best
predictor of brood survival. As with invertebrate counts,
broods located in areas with the highest median invertebrate
biomass (2.6 g/sample) were more than 7 times more likely to
have chicks at 2 weeks post-hatch (Sb ¼ 0.78, 90% CI 0.47–
0.94) than those with the lowest dry weight (0.1 g/sample;
Sb ¼ 0.10, 90% CI 0.02–0.41; Fig. 1). The 2 models with
invertebrate variables had a cumulative w of 0.87, indicating
Table 1. Results for binomial mixed effects modeling of factors affecting
brood survival at 2 weeks for Attwater’s prairie-chicken at 2 sites in the
Texas gulf coast prairie ecosystem during 2009–2013. K is the number of
parameters estimated, AIC is Akaike’s Information Criterion, DAIC is the
difference in AIC from the best model, and w is the model weight.
Model
K
AIC
DAIC
w
Sw
Invertebrate count
Invertebrate biomass
Brood locationsa
Year
Null
Site
3
3
3
6
2
3
81.85
85.46
87.06
88.05
88.99
89.67
0.00
3.61
5.21
6.20
7.14
7.83
0.75
0.12
0.06
0.03
0.02
0.01
0.75
0.87
0.93
0.96
0.99
1.00
a
Indicates the percentage of brood locations that were in a treated field
during the first 2 weeks.
901
Figure 1. Effect of (a) median invertebrate count, (b) median invertebrate
biomass, and (c) the percentage of brood observations in fields treated to
suppress red imported fire ants (RIFA) on probability of Attwater’s prairiechicken brood survival to 2 weeks. Gray area indicates 90% confidence
intervals.
considerably less support for the remaining 3 models
(Table 1). Although the percent of brood observations
within fire ant treated areas had substantially less support
than median invertebrate numbers, the b value (b ¼ 1.46,
90% CI 0.14–2.78) indicated that the estimated probability
of chicks being observed at 2 weeks was more than doubled
if a brood was always observed in a treated field (Sb ¼ 0.67,
90% CI 0.43–0.84) compared to a brood never observed in a
treated field (Sb ¼ 0.32, 90% CI 0.19–0.38; Fig. 1).
Impact of Fire Ant Suppression on Invertebrate
Abundance
We observed a reduction in fire ant activity as measured by
the fatty lure counts across all sites treated with Extinguish
Plus compared to reference sites; in 2011, fire ant counts on
treated sites averaged 53% lower than reference site counts
(range 35–68%) and in 2012 fire ant counts on treated sites
averaged 84% lower than reference site counts (range 64–
100%). Although invertebrate abundance and biomass were
highly variable across years and sites, 8 of 10 comparisons
(5 sites, 2 years) for each dataset showed positive treatment
responses (Fig. 2). The best supported models for the
total, mobile-flier, and ground-dwelling invertebrate counts
included year, site, an interaction between year and site,
902
and a treatment effect (Table 2). For mobile-fliers, the
model selection results for counts also indicated an
interaction between treatment and year. Residuals of top
models were approximately normally distributed with no
strong pattern of overdispersion or heteroscedasticity.
The confidence intervals of the relative treatment effect
did not overlap 1 in any of the top count models (Table 3).
For all invertebrates combined, count per plot for treated
fields was an estimated 1.27 (95% CI 1.16–1.40) times
higher than reference fields across all sites and years.
Mobile-flier and ground-dwelling invertebrates had similar
responses to treatment (Table 3), although there was
support for differences in mobile-flier treatment effects
by year.
For biomass, treatment was in the top models for grounddwelling invertebrates and total invertebrates with less
support for a treatment effect for mobile-flier invertebrates
(Table 2). The best transformations for total and grounddwelling invertebrate biomasses were the 0.20 and 0.24
power transformations, respectively. Residuals of the top
models were approximately normally distributed with no
strong pattern of overdispersion or heteroscedasticity. Total
invertebrate biomass had a nearly identical response to
treatment as total count; total biomass in the treatment fields
was an estimated 1.26 (95% CI 1.15–1.39) times higher than
the biomass in reference fields across all sites and years. For
mobile-flier biomass, there was evidence that fire ant
treatment slightly increased the probability that mobile-flier
invertebrates would be present at a plot in 2011 but not in
2012, and there was no evidence that fire ant treatment
increased the biomass of mobile-flier invertebrates at a plot,
given that they were present (Table 3). For ground-dwelling
invertebrate biomass, there was evidence of a treatment and
year interaction (Tables 2, 3). Contrary to our expectation,
there was no clear evidence of treatment effects increasing
during the second year of the study (Table 3); grounddwelling invertebrate biomass was the only 1 of the 6 analyses
to exhibit a larger treatment effect in the second year of
treatment.
DISCUSSION
Juvenile grouse (subfamily Tetraoninae) experience highest
mortality during the first 2 weeks of life (Hannon and Martin
2006). Data collected in this study clearly demonstrate that
availability of invertebrates during this period was a major
factor limiting survival for young Attwater’s prairie-chickens.
Our observations were consistent with other studies
affirming the importance of invertebrates to galliform chicks
during the first weeks of life, and for prairie-chicken species
in particular (Lehmann 1941, Jones 1963, Savory 1989,
Hagen et al. 2005). Lehmann (1941:26) observed that
stomachs of 3 juvenile Attwater’s prairie-chickens contained
88.5% insects. Population simulations suggest that to reach
Attwater’s prairie-chicken recovery goals, brood survival
must increase to levels comparable to stable greater prairiechicken (Tympanuchus cupido) populations and those of
historical wild Attwater’s prairie-chickens (Pratt 2010). Pratt
(2010) reported 69% survival of greater prairie-chicken
The Journal of Wildlife Management
79(6)
Figure 2. Median invertebrate counts and biomass/25 sweeps at 5 locations in Texas during 2011–2012 including Attwater Prairie Chicken National Wildlife
Refuge (APCNWR). The dark line indicates median value with boxes representing the 75% and 25% quantiles and the lines representing the 2.5% and 97.5%
quantiles.
broods to 2 weeks post-hatch. Our analyses suggest that
Attwater’s brood habitat would have to support invertebrate
populations resulting in approximately 190 invertebrates/25
sweeps to achieve this brood survival (Fig. 1). Invertebrate
abundances of this level or higher were observed only for 16%
of broods, suggesting that much of Attwater’s brood habitat
supported less than the desired levels of invertebrate
abundance. However, broods with all of their locations
within treated fields had survival estimates at 2 weeks of 67%
(90% CI 43–84%), suggesting that widespread fire ant
control could result in Attwater’s prairie-chicken brood
survival comparable to that observed for stable greater
prairie-chicken populations (Pratt 2010).
Data collected in this study also clearly demonstrate that
fire ants have caused a substantial reduction in invertebrate
abundance within historical and extant Attwater’s habitats,
despite considerable variation among sites and between years
(Fig. 2). Insects, especially immature forms, make up the
bulk of fire ant diets (Hays and Hays 1959). This suggests a
likely mechanism of fire ant impacts on invertebrates, and
likely explains the more consistent increases of grounddwelling invertebrates due to treatment for both count and
Table 2. Results for linear and generalized linear mixed effects modeling of the effect of fire ant treatment on invertebrate counts and biomass at 5 sites in
the Texas gulf coast prairie ecosystem during 2011–2012. K is the number of parameters estimated, DAIC is the difference in Akaike’s Information Criterion
(AIC) from the best model, and w is the model weight. Only models with DAIC <10 are shown.
Data seta
Count
Mobile-fliers
Ground-dwelling
Total
Biomass
Mobile-fliers (presence)
Mobile-fliers (if present)
Ground-dwelling
Total
a
b
Fixed effectsb
K
DAIC
w
Site þ year þ site year þ treatment þ treatment year
Site þ year þ site year þ treatment
Site þ year þ site year
Site þ year þ site year þ treatment
Site þ year þ site year þ treatment þ treatment year
Site þ year þ site year þ treatment
Site þ year þ site year þ treatment þ treatment year
14
13
12
13
14
13
14
0.0
2.3
9.6
0.0
1.7
0.0
2.0
0.75
0.24
0.01
0.70
0.30
0.73
0.27
Site þ year þ site year þ treatment þ treatment year
Site þ year þ site year þ treatment
Site þ year þ site year
Site þ year þ site year
Site þ year þ site year þ treatment
Site þ year þ site year þ treatment þ treatment year
Site þ year þ site year þ treatment
Site þ treatment
Site þ year þ site year þ treatment
Site þ year þ site year þ treatment þ treatment year
13
12
11
12
13
14
13
8
15
16
0.0
2.6
9.0
0.0
1.5
0.0
1.1
6.1
0.0
0.6
0.78
0.21
0.01
0.66
0.25
0.61
0.36
0.03
0.58
0.42
Mobile-fliers ¼ more mobile, flying invertebrates; ground-dwelling ¼ less mobile invertebrates; and total ¼ those 2 groups combined.
A random intercept of sampling plot was included in all models, except total biomass which had a random intercept of sampling plot and a random slope
across the 3 sampling periods.
Morrow et al.
Fire Ants and Attwater’s Prairie-Chicken
903
Table 3. Treatment effects (relative to untreated reference plots) of red
imported fire ant suppression on insect abundance across 5 sites within the
Texas gulf coast prairie during 2011–2012 as estimated from linear and
general linear mixed effects models.
Type
Data seta
Yearb
Count
Mobile-fliers
Biomass
Ground-dwelling
Total
Mobile-fliers presence
2011
2012
Both
Both
2011
2012
NAd
2011
2012
Both
Mobile-fliers if present
Ground-dwelling
Total
Treatment effectc
1.53
1.10
1.28
1.27
1.12
1.00
(1.22–1.91)
(0.82–1.50)
(1.15–1.43)
(1.16–1.40)
(1.00–1.17)
(0.86–1.14)
NA
1.19 (1.07–1.31)
1.34 (1.17–1.52)
1.26 (1.15–1.39)
Mobile-fliers ¼ more mobile, flying invertebrates; ground-dwelling ¼
less mobile invertebrates; and total ¼ those 2 groups combined.
b
If the model selection process supported the inclusion of a year treatment interaction, treatment effects were given by year.
c
The treatment effect is relative to the reference plots (e.g., a treatment
effect of 1.53 indicates treated plots had 1.53 the abundance or biomass
of the reference plots). A treatment effect of 1 indicates no treatment
difference, <1 or >1 indicates a decrease or increase in abundance or
biomass due to treatment, respectively. Numbers in parentheses indicate
95% confidence intervals.
d
No evidence of treatment effects.
a
biomass, and between years of our study, compared to
mobile-fliers (Table 3). Ground-dwelling invertebrates
consisted of immature Orthopterans as well as others we
hypothesized to be less mobile (e.g., leaf hoppers, beetles,
spiders) and therefore more vulnerable to Attwater’s and fire
ant predation.
A likely explanation for lack of treatment effects and
substantially higher invertebrate abundance at APCNWR in
2012 (Fig. 2) is that APCNWR was actively managed to
increase insect abundance in Attwater’s brood habitat for
several years prior to our study because Attwater’s are the
focus of management at this site. Management practices
directed at increasing invertebrate abundance have included
patch burning (e.g., Fuhlendorf and Engle 2001, Roper
2003, Fuhlendorf et al. 2006, Doxon 2009, Doxon et al.
2011) and removal of cattle during January–June. Fire breaks
on APCNWR are disked to encourage early successional
forbs, which often support high insect numbers (Jones 1963,
Green 1984, Hill 1985). Disking fire breaks also occurred at
the Midwell and West Clarkson but not at TCPP or Tivoli.
Removal of grazing during the first half of the growing
season, and reduction in overall grazing pressure due to
the 2011 drought, also likely explain at least part of the
anomalous responses at APCNWR compared to other sites.
Cattle are highly selective for forage based on palatability
and availability, and have had substantial impacts on spring
forbs (e.g., Coreopsis spp.) known to support high insect
numbers at APCNWR (M. E. Morrow, personal observation). Quinn and Walgenbach (1990) observed that grazed
sites supported higher numbers of obligate grass-feeding
grasshoppers, whereas ungrazed grasslands were dominated
by mixed forb- and grass-feeding grasshoppers, implying
that grazing may have reduced forb abundance and thus forbfeeding grasshoppers in that study.
904
Other factors, including the ecology of individual species,
affect invertebrate species abundance within a grassland
ecosystem (Quinn and Walgenbach 1990, Panzer 2002,
Debano 2006). Weather, especially temperature and rainfall,
are particularly important (Uvarov 1977). The 2 years of our
study were very different, with drought conditions prevailing
during 2011 but not during 2012. This likely explains why
year was included in all of the top models in the linear and
generalized linear mixed effects models of invertebrate
abundance (Table 2). It may also explain the smaller
treatment effects we observed (1.27 and 1.26 reference
plots for invertebrate numbers and biomass, respectively)
across all sites compared to other studies (Porter and
Savignano 1990, Allen et al. 2001). Much of the increased
invertebrate abundance for APCNWR in 2012 (Fig. 2),
irrespective of treatment, was attributable to large numbers
of a very small, winged Psyllid (Psyllidae). It is likely that
environmental conditions were favorable in 2012 for this
species at APCNWR and populations exploded, overwhelming any treatment effect. A similar phenomenon with
grasshoppers occurred at APCNWR in 2010 (APCNWR,
unpublished data) and at some locations in Goliad County in
2012 (J. Kelso, The Nature Conservancy of Texas, personal
communication). Although undoubtedly fire ants have an
overall suppressive effect on insect abundance during the
May–mid-June Attwater’s brooding period, there are likely
times when environmental conditions are favorable for
eruptive invertebrate population growth that are able to
overcome the suppressive impacts of fire ants. Studies of
longer duration than 2 years would be necessary to evaluate
that hypothesis. This high level of spatial and temporal
variation in invertebrate densities may result in a lack of an
observed treatment effect of fire ant suppression in some
years or locations.
Reports of adverse direct impacts of fire ants to native fauna
are numerous (Allen et al. 2004), but to our knowledge few
studies address indirect impacts of this invasive species.
Indirect effects related to a reduction in invertebrates due to
fire ants have been hypothesized for diverse groups of species
including northern bobwhites (Colinus virginianus; Allen
et al. 1995), loggerhead shrikes (Lanius ludovicianus; Lymn
and Temple 1991), Texas horned lizards (Phrynosoma
cornutum; Allen et al. 2004), and the endangered Houston
toad (Bufo houstonensis; USFWS 2009). Allen et al. (2001)
observed that both insect biomass and loggerhead shrike
abundance were negatively correlated with fire ant abundance.
MANAGEMENT IMPLICATIONS
Our study suggests that for Attwater’s recovery to be
successful, fire ants must be controlled or other methods
found to mitigate their effects such that invertebrates used as
food by Attwater’s chicks are present in sufficient numbers to
support high brood survival. Given the Attwater’s high
potential reproductive rate (Pratt 2010, USFWS 2010),
improving brood success should result in substantial progress
toward recovery of this critically endangered species. Results
of our study indicate that the ecological disruption caused by
The Journal of Wildlife Management
79(6)
invasive fire ants goes beyond direct impacts to individual
species, and may contribute to a trophic cascade affecting
other species occurring throughout the acquired range of the
red imported fire ant. Indeed, long-term population declines
observed for northern bobwhites and wild turkeys (Meleagris
gallopavo; North American Bird Conservation Initiative
2009, Nickens 2013) may be partially attributable to indirect
impacts of fire ants, as chicks of both species depend on
insect-rich diets (Palmer et al. 2001, Randel et al. 2007). We
strongly suggest that additional research is needed to fully
elucidate the indirect impacts of fire ants on other species and
to provide guidance for corrective management actions.
ACKNOWLEDGMENTS
We gratefully acknowledge support by the National Fish and
Wildlife Foundation (Project Number: 2010-0008-000/
509), Central Life Sciences, Society of Tympanuchus
Cupido Pinnatus (STCP), U.S. Fish and Wildlife Service,
Attwater Prairie Chicken National Wildlife Refuge, Texas
A&M University Department of Entomology, Texas
AgriLife Extension Service, and The Nature Conservancy
of Texas. A. Calixto, Texas AgriLife Extension Service,
provided technical guidance on study design, and helped
collect data during 2011. We are especially grateful to private
landowners who allowed access to their properties, and to C.
Carter who served as the liaison with those landowners.
Numerous volunteers and Student Conservation Association
interns at APCNWR, and research technicians at Texas
A&M University, Department of Entomology, assisted with
field work and sample processing. D. Mueller was notably
diligent. A. Pratt (STCP) and J. Kelso, The Nature
Conservancy, collected brood data at Goliad County study
sites. J. Mueller provided comments on an earlier manuscript
draft. The findings and conclusions in this report are those of
the authors and do not necessarily represent the views of
the U.S. Fish and Wildlife Service. Use of trade names does
not constitute endorsement by the authors.
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